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University of Glasgow Institute of Biodiversity, Animal Health & Comparative Medicine

Discussion Papers in Environmental and One Health Economics

Paper Number 2019 – 01

The economic benefits of invasive species management

Nick Hanley (Institute of Biodiversity, Animal Health and Comparative Medicine,

University Of Glasgow)

and

Michaela Roberts (James Hutton Institute, Aberdeen)

Abstract Invasive species are known to cause significant negative impacts to ecosystems and to people. In this paper, we outline the range of approaches to estimating the economic costs of invasive species, and thus the benefits of management programmes. We provide examples of the application of these approaches to specific instances of invasive species management, and conclude by asking what economists need from ecologists to implement these approaches.

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1. Introduction. Invasive species are those introduced to a novel environment with negative ecological, economic or social impacts (Mooney, 2001). These negative impacts have been increasingly recognised in both the ecological and economic literatures, as awareness of the impacts of invasive species grows, and as globalisation increases the pathways and speed of invasions (Seebens et al., 2018; Smith et al., 2018). In this paper, we outline the approaches that economists take to measuring the costs of invasive species, taking in both commercially-valued losses and “non-market” effects, whilst noting that economic benefits arise from non-native species in many instances. We discuss some of the problems of applying economic valuation approaches to invasives, and review what information from ecologists is needed for such methods to be used. Our understanding of the impacts of invasive species is greatest for those that have impacts on agriculture and forestry (Vilà et al., 2010).However the overall impacts of invasive species are wide ranging across ecosystems (Pejchar & Mooney, 2009). Invasive species damage food production (Engeman et al., 2010) and can act as disease vectors (Medlock & Leach, 2015). Due to the complex nature of ecosystems it is likely that we do not yet understand the full impacts. Negative impacts mean that controlling invasive species is becoming increasingly important for society. However, control often incurs high costs (Martins et al., 2006), and may be met with social opposition (Roberts, Cresswell, & Hanley, 2018; Sheremet, Healey, Quine, & Hanley, 2017), particularly where invasive species have acquired cultural values (Roberts et al., 2018). When control efforts are unsuccessful, and/or where the damages associated with the invasion are low relative to the costs of control, then it may be socially desirable to abandon control measures and instead manage the impacts of damage (Rolfe & Windle, 2014). The economic benefits of invasive species management are equal to the avoided costs of damages from invasives were control not to be implemented. The size of benefit thus depends on the speed of spread for new invaders, or area of invasion for established populations; the damages per “unit” (e.g. per possum, per infected km2); and how many people are affected by these damages. The net benefits of a management programme are equal to the value of avoided damage minus the costs of control and any benefits forgone that were previously provided by the invasive species (e.g. fuelwood production or hunting opportunities). Net benefits of interventions therefore also depend on the effectiveness of control options, when and where such options are implemented, and for how long they are implemented, since this will determine the time period over which (discounted) benefits and costs of control are added up (Figure 1). As this paper focuses only on the economic benefits of invasive species control, there are many issues which economists have contributed to on this subject which we do not cover. We do not give much attention to the costs of control, and no attention at all to the economically-optimal level of management effort, the timing of control actions (Sims & Finnoff, 2013), or the use of economics in modelling invasive species spread. For an excellent overview of many of these issues, see Epanchin-Niell (2017). We also do not consider the

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question of how much resources to invest in biosecurity measures to try to prevent potentially-invasive species from entering a country.

Figure 1 Calculation of net benefit of invasive species management programmes. A full economic evaluation of control options would discount the stream of predicted benefits and costs of control over time, using an appropriate social rate of discount (see Hanley and Barbier, 2009).

The objective of this paper is to illustrate the economic considerations which should be incorporated into invasive species control or management, and provide examples of the methods which could be applied. We do not intend to provide a comprehensive road map for economic valuation, and would direct readers looking for further guidance to Hanley and Barbier (2009) for a more in depth but accessible introduction. In what follows, we give an overview of the nature of the control problem, the kinds of economic benefit which might be gained from management of invasives, and how such benefits can be measured. Examples are provided from recent estimates of the benefits of control, focussing on what we will define below as “non-market” benefits. 2. A growing problem? The total number of invasive species is increasing worldwide (Huang, Haack, & Zhang, 2011; Seebens et al., 2017, 2018; Smith et al., 2018). The number of new invasions, as well as the number of individual species recognised as invasive, has increased steadily since 1800, with an increased rate of introduction after 1950 (Seebens et al., 2017, 2018). The rise is linked to the expansion of global trade, specialisation in production and increased connections to previously isolated locations (Seebens et al., 2018). Climate change also opens up new pathways for introduction and for range expansion of already-introduced species. In China, USA and UK, the number of invertebrate pests has increased with rising mean temperatures, even after accounting for increased trade (Huang et al., 2011), while increased temperatures in Europe have led to the establishment of mosquito species and associated vector-borne diseases (Medlock & Leach, 2015). Though the trends for increasing numbers of invasive species seem widespread, it is important to account for increased effort in identifying invasive species which increases the probability of detection (Costello & Solow, 2003). As invasive species increase, so do actions to control them. Costs of invasive species control are poorly reported, often contained within grey literature and limited in the species

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considered and geographical scope (Brooke, Hilton, & Martins, 2007; Holmes et al., 2015; Martins et al., 2006). Although cost reporting is scarce, those projects that do report costs generally show increased cost effectiveness over time, particularly where actions occur in the same geographic location (Carrion, Donlan, Campbell, Lavoie, & Cruz, 2011; Donlan & Wilcox, 2007; Martins et al., 2006). Despite increased efficiencies, the total costs of invasive species control are most likely increasing. This is largely due to control actions taking place in increasingly complex locations, namely locations with multiple invasive species (Glen et al., 2013). As the number of invasive species and the complexities of control increase, so does the importance of predicting the costs for prioritisation of control actions. The most consistent predictor of costs is the size of control area, with larger areas having smaller per ha costs (Martins et al., 2006), however this is not consistent across all studies (Holmes et al., 2015). Species type can also have an impact, as can remoteness (Holmes et al., 2015; Martins et al., 2006). Prioritisation of invasive species control needs to account for the costs of control, as well as the damages avoided, and any values of the invasive species itself (Cook, Thomas, Cunningham, Anderson, & De Barro, 2007; Donlan, Luque, & Wilcox, 2015; Roberts et al., 2018). 3. Typology of economic values from invasive species management The impacts of invasive species are associated with a range of costs and benefits, with many species having both positive and negative values depending on context (Goodenough, 2010; Oreska & Aldridge, 2011; Shackleton, Shackleton, & Kull, 2018). Impacts vary through space and through time. For example, established brown trout populations worldwide have less current ecological impact, from a population to an ecosystem level, than newly arrived populations (Závorka, Buoro, & Cucherousset, 2018). Though it is widely recognised that invasive species have impacts on wellbeing, our understanding of these impacts from an ecological and economic point of view is very incomplete (Vilà et al., 2010). In this section, we use an economics perspective to categorise the types of impacts arising from invasive species, organising these into direct and indirect, positive and negative, and market and non-market impacts. In all cases, we define an economic benefit or cost in terms of an impact on human well-being which can be attributed to a change in the population or spatial distribution of an invasive species. Table 1 summarises this classification system.

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Table 1 types of economic cost or benefit due to invasive species.

Classification Example Are the example impacts valued by the market?

Direct: positive impact on human well-being

Increased recreational hunting opportunities due to introduction of deer in New Zealand

Partially, if land managers charge hunting fees.

Direct: negative impact on human well-being

Spread of mosquito-borne human diseases in Europe due to climate change

Not where health care services are publicly provided

Indirect: positive impact on human well-being

Use of non-native species (eg sitka spruce) in UK timber production

Yes: timber markets reflect the value of yield increases from using non-native species

Indirect: negative impact on human well-being

Effects of ash die back on forest recreation and ecological quality of ash woodlands in UK

No: lost recreation benefits and deteriorating ecological quality are not priced by the market.

Direct impacts are those effects of invasives that directly impinge on the well-being of economic agents (consumers, taxpayers, producers). Indirect effects happen via a transmission mechanism. For example, Ash die-back is a disease of Ash trees caused by an invasive fungus Hymenoscyphus faxineus which was introduced to the UK from continental Europe in 2012. The negative effects on UK households of the loss of treasured ash woodlands are brought about via the effect of the fungus on ash trees. Both direct and indirect impacts can be thought of in terms of market-valued impacts and “non-market” values: we explain this distinction below. Invasive species can have a direct, positive impact on communities through provision of resources (Shackleton et al., 2018). Common Pheasants Phasianus colchicus in the UK countryside provide a source of enjoyment for hunters as well as income for the estates which manage them. The species was introduced to the UK from China as a game species in the 18th century (Statistical Accounts of Scotland, 1795). Large numbers of pheasants are now bred and released each year on UK sporting estates (Robertson, 1996). Rhododendron ponticum was first introduced to the UK as an ornamental plant, and has value for gardeners due to its easy care, evergreen nature, and bright flowers. Though attitudes towards R. ponticum have changed in recent years due to its fast spread and ability to outcompete native plants, removal projects are often met with public resistance (Williamson, 2006). Invasive species can also have direct, negative impacts. As well as being a game bird, pheasants are vectors of Lyme disease. Because pheasants remain infectious for up to ten weeks, the species acts as amplifiers for the disease in the environment, increasing risk of transmission to humans (Kurtenbach, Carey, Hoodless, Nuttall, & Randolph, 1998). The impact of invasive species as vectors of disease is predicted to increase as a result of climate change. Invasive mosquitos have increased in mainland Europe, bringing with them chikungunya and dengue fever, with cases reported in Italy and France. Climate models predict that climate change will lead to increased potential for these species to invade the UK (Medlock & Leach, 2015).

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Invasive species may have indirect impacts through altering resource provision, many of which effects can be valued using market prices. The most commonly studied impacts of invasive species are those on agriculture and forestry. Early back-of-the-envelope calculations for US agriculture show losses due to the effects of non-native weeds, pests and pathogens of around $65 million (Pimentel, Zuniga, & Morrison, 2005). , while invasive pests and pathogens such as emerald ash borer and Phytophthora ramorum cause damage to commercial forestry (Freer-Smith & Webber, 2017). Invasive species can also disrupt services: in the UK, rail services spend money to control R. ponticum growing near to railway lines to prevent service disruption (Williamson, 2006). Zebra mussels growing on boat hulls reduce fuel efficiency and can impact water filtration, thus effecting shipping costs (Oreska & Aldridge, 2011). Indirect impacts may also occur to non-market resources: wild species and landscapes whose economic value is not fully reflected in market prices. Among the positive ,indirect impacts of invasive species is the provision of food resources for a native species (Goodenough, 2010). In Davis, California, 13 native butterfly species exclusively use invasive species as breeding hosts, and therefore food for their young. Although these species exploit native food species elsewhere, the butterfly species would be locally extinct if their invasive hosts were controlled within Davis (Shapiro, 2002). Invasive species can also disperse seeds of native plants (Dungan, O’Cain, Lopez, & Norton, 2002; Goodenough, 2010). Indirect, non-market negative impacts of invasive species can also arise through adverse impacts on native ecosystems or species. Bird, reptile and mammal populations worldwide are well known to be decimated by domestic cats (Woods, Mcdonald, & Har Ris, 2003). In Guam, the brown tree snake is credited with the devastation of native bird species due to nest predation (Pimentel et al., 2001). Pathogens can also directly reduce populations. As noted above, Ash dieback as a result of the introduced fungi Hymenoscyphus fraxineus has led to large reductions in the ash population within the UK (Freer-Smith & Webber, 2017), which leads to losses in well-being for people who enjoy walking in ash woodlands. The loss of ash trees also represents a loss of pollution sinks, linked with increases in human cardiovascular and respiratory mortality (Jones & McDermott, 2017). Loss of ash trees due to the Emerald Ash Borer in the USA has been associated with a reduction in life satisfaction due to the loss of locally-valued woodlands (Jones, 2017). Indirect damage to native species can occur due to increased competition for food resources (Robertson, 1996). Because invasive species are often prolific within the introduced system, they can change native ecosystem functioning. The introduction of purple loosestrife and water hyacinth structurally changes wetlands in Europe and North America (Pimentel et al., 2001). These changes can have significant impacts on people who care about the ecological quality of waterbodies, or whose recreational experiences are diminished due to the adverse impacts of such invasives on swimming or boating opportunities. Again, these adverse impacts are signalled by market in a transparent manner. 4. Methods for valuing impacts Markets are good at signalling the value of a large set of goods and services which might be affected, directly or indirectly, by an invasive species. Market prices reflect the interaction of

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supply and demand, for example for timber, wheat or farmland. Supply curves provide information on the incremental cost of producing goods (e.g. timber from a forest), and how this cost varies across producers. Demand curves show how much buyers (consumers or other firms) are willing to pay for goods and services: that is, the value that they place on having such goods and services made available to them. Thus the demand curve for timber shows how much potential buyers are willing to pay for this good, which reflects the value of the good to them. Market forces mean that prices move to equilibrate demand and supply, so that the market price at any point in time shows both the incremental cost of producing the good and its incremental value to buyers. This means that when markets work well, market prices provide valuable signals on the social costs and benefits of changes in output. Thus, a good estimate of the economic costs of a 10% loss of agricultural output due to an invasive pest is the market value (price times quantity) of this change in output, less the costs of producing this crop. Note, however, that actually establishing and then identifying the size of the causal link between the introduction of a pest/pathogen and the effects on agricultural profits may be challenging. Moreover, farmers may change their management in response to an invasive species, so such simple calculations can give a misleading estimate of the true social costs of lost output. We need to factor in any such behavioural responses to changes in production risk in measuring economic damages. However, market prices do not send good signals about social costs and benefits when demand and supply curves do not reflect all of the costs associated with producing a good (for example, when dairy farming leads to increased water pollution) and/or all of the benefits of producing the good (for instance, when people use forests for recreation as well as for timber production). Such circumstances extend to cases where markets are simply missing for certain benefits and costs, such as the impacts of the invasive species on native biodiversity. Economists refer to both of these instances as a case of “market failure”, and for more than 100 years have known that market failure implies that market prices no longer provide adequate information on the social (economic) costs or benefits of changes in output. Market failure characterises many of the situations in which invasive species generate direct or indirect effects on human well-being. The spread of mosquito vectors of dengue fever brings about increases in morbidity which typically are not valued by markets: extra cases of dengue fever in Spain over the next 10 years can be partly valued though increases in treatment costs to the Spanish health services, but generate wider effects of people’s well-being which such approaches under-value. The introduction of stoats to New Zealand in the 1880s (to control rabbits) brought about significant and on-going predation losses to ground- nesting birds such as the kakapo and takahe (King, 1984). The introduction of possums Trichosurus vulpecula to New Zealand from Australia has led to increased bovine TB outbreaks in dairy cattle (the value of which is recognised by markets for cattle and for dairy products) and to ecological damages to native forest species such as rata, which go completely un-valued by the market (Department of Conservation, 2008). In the next section we explain how such non-market values can be measured. 4.1 Tools for non-market valuation Since the mid-1960s, economists have built up a tool kit of methods for estimating non-market values attached to the environment. Initially developed in the context of national park

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planning, water quality enhancements and public forest management, these methods have now expanded to be able to value changes in a very wide range of environmental benefits and costs, from changes in urban air quality to the conservation of wetlands. All of these methods can be used to value the non-market impacts of invasive species. Non-market valuation methods are usually categorised into 3 types (Hanley & Barbier, 2009):

- Stated preference approaches

- Revealed preference approaches

- Production function methods

All are based on the notion of maximum willingness to pay (WTP) as a standard measure of the economic value of a good to individuals, since in economics the value someone places on any good or service depends not just on their preferences, but also on how much they are willing to give up to obtain it.

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Table 2 summarises the applicability of each of the methods described below to valuing the non-market economic damages or benefits associated with invasive species. Stated and revealed preference approaches estimate the direct effects of environmental change on individual well-being. Stated preference methods ask individuals to make choices between different “attributes” and the cost of providing a good. In this way, people show the value they place on, for example, avoiding damage by possums to native forests, or funding a fire ant control programme. Two stated preference methods dominate the literature: contingent valuation and choice modelling (Hanley & Czajkowski, 2018). In contingent valuation, people vote on whether they agree with a specific change in the provision of an environmental good (e.g. delaying invasive species arrival, maintaining human health and recreation in US waterbodies) at a specific cost to them (e.g. a payment of $48 to delay arrival for one year, McIntosh, Shogren, & Finnoff, 2010). In choice modelling, people make choices between different “bundles” of environmental goods – such as different measures for invasive species control – as a function of the attributes of this good (e.g. forest ownership, type of forest, control action) where one of these attributes comprises a cost of providing the good (e.g. an increase in local taxes). One of the attributes of the good over which people make choices could be populations or spread of an invasive species, or the impacts of a species on, for instance, forest quality. Attributes could be also used to describe the different potential components of a management plan, such as whether biocides are used (Sheremet et al., 2017). Stated preference methods have the disadvantage that they are not based on actual payment for the good. However, they offer many advantages: widespread applicability and the ability to measure both non-use and use values (Hanley & Barbier, 2009). Considerable effort has been devoted to understanding how best to design such studies, how to minimise the extent of hypothetical market bias, and what kinds of econometric model are most appropriate given the nature of the data and the range of information-processing and choice processes that individuals may employ in responding. For example, the Sheremet et al (2017) paper noted above focusses on the issue of heterogeneity in peoples’ preferences for invasive control strategies in UK woodlands when analysing the stated choice data.

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Table 2 methods for valuing non-market impacts of invasive species

Method Type of impact valued

Advantages Disadvantages Examples

Stated preference: Choice experiment

Non-use; Recreation; Landscape; Biodiversity; Cultural heritage

Wide range of applications; Hypothetical markets means can be used to value planned control measures or estimate costs of potential invasive species. Can value multiple attributes and compare preferences.

Hypothetical markets not based in real payments. High cognitive burden for participants.

Preferences for control of invasive tree diseases (Sheremet et al., 2017)

Stated preference: Contingent valuation

Non-use; Recreation; Landscape; Biodiversity; Cultural heritage

Wide range of applications. Hypothetical markets means can be used to value planned control measures or estimate costs of potential invasive species. Lower cognitive burden for participants than choice experiment.

Hypothetical markets not based in real payments. Good can (typically) only be valued as a whole, not by individual attributes.

Values of delayed arrival date of invasive species (McIntosh et al., 2010)

Revealed preference: Travel cost models

Recreation Based in real behaviour. Where multiple alternative sites exist can value individual attributes.

Limited number of scenarios in which this can be applied. Relies on existing markets and presence of alternative sites. Underestimate of value as cannot account for non-use values.

Recreation impacted by feral herbivore grazing (Peh et al., 2015)

Revealed preference:

Housing markets Based in real behaviour.

Limited scenarios in which method

Invasive aquatic

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Hedonic pricing

Potential to value individual attributes.

can be applied, very rarely used in invasive species research. Underestimate of value as cannot account for non-use values.

plants impact on waterfront house prices (Zhang & Boyle, 2010)

Production function

Crop production; Production forests; Livestock; Human health

Based in market values.

Limited scenarios in which applicable. Cannot account for many direct effects on utility.

Impact of ragweed on costs of allergies (Richter et al., 2013)

Revealed preference methods are based on actual behaviour rather than stated choices. Travel cost models use people’s expenditures on outdoor recreation trips (e.g. mountaineering day trips, fishing trips, bird watching visits) to infer the demand for the natural resources (mountains, rivers, wetlands) which are the destinations of these trips. More relevantly, if there is a quality change at a given site (e.g. a loss of tree cover due to an invasive pest) or if a site is no longer available (e.g. if a suburban forest site is closed to recreationalists because of the presence of oak processionary moth caterpillars), then the economic losses of this closure of a site or due to a decline in site quality can be inferred. Similarly, the economic benefits of an increase in deer numbers which allows hunters in New Zealand to “consume” more days of hunting recreation could be valued using this approach. The second revealed preference method is the hedonic pricing approach. This examines the role of spatial and temporal variations in environmental quality on house prices, based on the assumption that people are willing to spend extra on a house, all else being equal, to “buy” better local environmental quality. Thus, houses closer to better quality urban green spaces, or with better air quality, or lower noise levels, will, on average, attract higher bids from house buyers than properties further away from green space, or with higher pollution levels, or which are in noisier neighbourhoods. The method can be used to value the economic benefits of managing invasive species which impact house prices: examples include the effects of non-native aquatic plants on house values in the US (Horsch & Lewis, 2009; Zhang & Boyle, 2010). However, the class of invasives for which this method is applicable will be rather limited. Production function methods link invasive species population changes to impacts on

commercial crops and livestock, or to human health outcomes. Figure 2 shows some of the

possible linkages. The right and left arrows show epidemiological models which translate

the change in the “arrival” of an invasive pathogen, for instance, into its effects on

commercially-grown crops. Crop losses can be valued using market prices. For forests, the

arrival and spread of the pathogen may change the optimal management of the forest in

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terms of the optimal rotation period and/or the optimal planting mix of species

(MacPherson et al, 2017, 2018). Moreover, if we think about the potential irreversibility of

certain invasive species control options (e.g. the introduction of a natural predator), and the

likelihood that we will learn more about the epidemiology and impacts of the invasive over

time, then real options models can be used to estimate the costs of acting too soon or too

late (Sims & Finnoff, 2013). For human health effects – for example, in terms of cases of

dengue fever in a country due to the arrival of the mosquito Aedes aegypti which spreads

the dengue virus – several valuation methods exist, including the use of stated preference

methods to measure WTP for reducing disease risks, and the Costs of Illness approach,

which sums medical system care costs and lost earnings due to sickness. Here, an

epidemiological model links changes in the invasive species to changes in human health

status; and then an economic valuation is placed on this change in health status.

Figure 2 Link between invasive species population changes to impacts on commercial crops and livestock, or to human health outcomes, as used in production function approach.

5. Examples of non-market valuation associated with invasive species impacts In this section, we provide some examples of applications of the methods described above to estimating the benefits of invasive species control. For reasons of space, we focus on the non-market benefits of control programmes. As explained above, choice modelling asks respondents to make choices between different bundles of (environmental) attributes, allowing the researcher to infer the economic value which people place on each of these attributes (Hoyos, 2010). The method has been applied to uncover aspects of a control programme most valued by citizens, such as the spatial targeting of control measures against invasive fire ants in Queensland (Rolfe & Windle, 2014). An example of a choice card used in the study by Sheremet et al (2017) on public WTP for programmes to counter invasive forest pests and pathogens is included in the SI. The choice modelling method has been extensively used in a wide range of environmental and conservation management contexts (e.g. Roberts, Hanley, & Cresswell, 2017). Chakir et al (2016) apply the approach to quantify the impacts of the invasive Asian ladybird species Harmonia axyridis (the Asian ladybird) on French citizens. The Asian ladybird was deliberately introduced to France as a bio-control measure for aphid management in agriculture in the 1990s. It has since spread rapidly, and is associated with undesired negative impacts on native biodiversity, housing (due to overwintering in large numbers) and wine production (due to

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tainting). However, the presence of the Asian ladybird allows farmers to use lower volumes of pesticides for aphid control. Chakir et al included the following attributes in their experimental design, to represent the environmental management problem at hand (levels of the attributes are shown in parentheses):

● level of pesticides used in agriculture (status quo, a 3% increase over 5 years, a 3%

decrease, according to populations of Asian ladybird present)

● population level of the native 2-spotted ladybird Adalia bipunctata which is

adversely affected by the Asian ladybird (levels: not present in France; rare;

abundant)

● damages to humans due to presence of Asian ladybirds overwintering in houses

(defined as % of housing affected varying from 1% to 15%)

● cost of an Asian ladybird control programme to the French taxpayer.

Results showed that across the 464 respondents who completed all of the choice tasks, people were willing to pay to protect the native 2-spot ladybird and to reduce nuisance to householders; but they were also willing to pay to reduce pesticide use. This means that the French population would value a programme to protect/restore native biodiversity by reducing Asian ladybird populations, but would require compensation to make up for any increase in pesticide use for aphid control that this made necessary. Interestingly, WTP to remove the negative effects of Asian ladybirds was higher than the compensation needed to offset increases in pesticides. The results also show support for public research programmes into alternative ways of controlling this invasive species (Chakir, David, Gozlan, & Sangare, 2016). An example of a contingent valuation study of the benefits of invasive species control is McIntosh et al (2010). Costly control measures may only delay the arrival of an invasive species in an area, rather than guarantee it will never arrive. A nation-wide survey of US households elicited their maximum WTP to delay the arrival of aquatic invasives such as fish (common carp), molluscs (zebra mussels), crustaceans (rusty crayfish) or water plants (Purple loosestrife) to inland water bodies in “regional” lakes and rivers in the USA, defined as places to which the respondent could drive in no more than 2 hours. Scenarios presented included delaying invasions by one year or 10 years for high or low levels of impact. Impacts were described in terms of effects on human health, the economy, recreation and navigation. Results showed that mean WTP to delay impacts by 10 years was five times greater than that to delay impacts by only one year. The main policy conclusion was that even short delays in arrival could generate significant economic benefits (around $4-$5.5 billion). Another contingent valuation study is reported in Meldrum et al (2013). The authors estimate the non-market benefits of managing white pine blister rust in US forests. This invasive fungus has caused significant ecological damages to high-elevation forests in Western North America and Quebec. The authors explore how respondent attitudes to forest protection and use affect their WTP to reduce the spread of this pathogen. Responses from a random sample of Western US households were used to elicit mean WTP estimates for a white pine blister rust control programme across a varying part (30% to 70%) of 2 million acres of high-altitude pine forest. Just under half of respondents had a positive WTP for the programme. Mean WTP for

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the protection programme was around $300 across all of those respondents with a positive WTP. The study shows how respondents’ attitudes to why an invasive species control programme should be implemented can be incorporated into models of the estimated benefits of such a programme (Meldrum, Champ, & Bond, 2013). It is more difficult to apply travel cost models than stated preference approaches to invasive species management, since one has to find a way of specifying a quantitative relationship between the abundance or spatial distribution of the invasive species, and recreational site quality. Then, one needs to find a relationship between the number of visits individuals make to the recreational site and this invasive-dependent site quality index. Perhaps due to these requirements, it is hard to find examples in the literature of fully-developed travel cost model applications to the benefits of invasive species control. One partial analysis is presented in Peh et al (2015). The authors study the effects of feral goats and pigs on ecosystem quality in the Centre Hills, Montserrat, and try to quantify the economic benefits of on-going management of these feral species. Feral pigs and goats have adverse impacts on forest understorey through grazing which affects endemic and rare bird species such as the Montserrat oriole Icterus ober which facilitates the rapid spread of invasive non-native plants such as guava Psidium guajava. The authors consider three benefits of feral livestock control: enhanced carbon storage, nature-based tourism, and hunting. For nature-based tourism, a travel cost analysis was undertaken, based on interviews with overseas visitors. Spending on accommodation, meals and car rental was used as a measure of the total expenditure by overseas visitors. Sampled visitors were asked how likely it was that they would cease visiting if ecological quality in the Centre Hills declined (specifically, if “..the unique animals of Montserrat have disappeared”). These responses were used to produce an estimate of lost tourism values if feral livestock control were abandoned. However, as the authors note, this was an incomplete application of the travel cost approach, since only parts of travel costs were measured, whilst no allowance was given to tourists being willing to pay more than their actual travel costs to visit the reserve. For a similar application of the travel cost approach to the spread of rusty crayfish in Wisconsin, see Keller, Frang, & Lodge (2007). Hedonic price applications to measuring the benefits of invasive species control are, by definition, limited to instances where the invasive species affects house prices, through changes in the desirability of living in a specific location. For instance, the spread of an invasive aquatic plant can change the benefits of living at a lakeside location if this means that recreational opportunities are reduced. Zhang and Boyle (2010) use the hedonic price approach to study the relationship between house prices at lakeside locations in Vermont and the spread of Eurasian watermilfoil Myriophyllum spicatum. This plant spreads rapidly, crowding out native water plants and reducing recreational opportunities (swimming, fishing) as it forms dense mats. What is especially interesting about Zhang and Boyle’s work is that they make use of two indexes for water quality: one based solely on the abundance of milfoil, and one based on the abundance of all water plants including milfoil. The paper is also unusual in being able to make use of property-specific values for milfoil and total aquatic plant abundance, rather than more spatially-aggregated measures. Two alternative functional forms are used to represent the possible effects of milfoil abundance on house prices, quadratic and exponential. Results showed that whilst milfoil abundance on its own had no significant effects on house prices, total aquatic plant abundance (including milfoil) did. Marginal effects of increasing total aquatic plant coverage along a 6 point scale were

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computed, showing that, for example, reducing the coverage of aquatic plants from the highest level (6) to the next lowest level (5) would increase average house prices by around 20%. Production function approaches estimate costs of invasive species through their impacts on an associated market (eg the market for an agricultural crop, when the invasive species adversely affects farm outputs), or on human health outcomes. Invasive plants can cause allergic reactions, having economic costs due to healthcare and productivity losses from days of work lost. Richter et al (2013) estimated the costs of allergic reactions arising from ragweed throughout Austria and Bavaria, modelling costs under alternative climate scenarios for the period 2011-2050. To estimate the population affected by ragweed the authors first model the spread of ragweed under different climate scenarios. Under current climatic conditions, the mean annual cost to Austria and Bavaria of ragweed, in terms of health and productivity losses, is estimated to be €291 million Euro per year between 2011-2050, rising to €333-365 million per year under climate change scenarios. Because the cost estimates from this study are linked to modelled expansion of ragweed range, the models also predict that costs will increase over time, rising from €133 million in 2005 to €422 million in 2050 under current climate. 6. Values associated with how invasive species are managed. Multiple options are often available to managers to respond to an invasive species threat. Examples of “damage reduction activities” include lethal controls (e.g. poisoning), the fencing of vulnerable habitats, and felling or spraying of invasive pests and pathogens which affect forests. The public may well have preferences over these control options which should be taken into account in any Cost-Benefit Analysis of control measures. That is, if people would rather animals were not controlled by poisoning, then their WTP to avoid this measure being implemented is a non-market cost of the damage reduction activity which should be added to the financial costs of shooting-based programmes. Even if a full Cost-Benefit Analysis is not undertaken, one can take the perspective that damage reduction activities should be chosen with a view to somehow balancing their social acceptability with their ecological effectiveness and financial cost (Roberts et al., 2018). There is evidence that the public care about which invasive control measures are used or proposed (Bremner and Park, 2007). Lethal control is an obvious example. Hanley et al (2003) used to choice experiment to show that the willingness to pay for a goose management programme on the island of Islay (Scotland) was significantly reduced if the control programme included shooting. This result was found for the Scottish general public and for visitors to the island. Interestingly the WTP of residents of Islay was not significantly reduced by the use of shooting (Hanley, MacMillan, Patterson, & Wright, 2003). For forest disease control, Sheremet et al (2017) found that the UK public had a negative WTP for control options which consisted of either clear-felling infected forests, or which made use of chemical or biocide spraying. Jepson and Akakelyan (2017) found a negative attitude to the use of GM breeding methods as a response to ash die-back, and that this attitude varied significantly according to respondent’s age and education, and according to where GM-modified ash trees were planted in the landscape (Jepson & Arakelyan, 2017). Finally, Fleischer et al (2013) used choice modelling to study the preferences of Israeli citizens for different control options

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designed to respond to invasion by the Dwarf honey bee Apis florea. The attributes used in the design were impacts on two native plant species (Calotropic procera and Lupinus pilosus); the nature of a pesticide-based control programme targeted at the dwarf honey bee; and donations to a fund to pay for the programme. People were willing to pay between US$6 - $17 per month for a control programme, but this declined for around 25% of the sample when a pesticide was used (Fleischer, Shafir, & Mandelik, 2013). 7. Discussion

In the preceding sections, we saw how a number of different tools are available for estimating the economic benefits of invasive species management. In this final section, a short discussion is provided on some of the main challenges in applying these methods in this specific context, again focussing on non-market impacts. The first problem to note turns on the issue of how much we require people to know about an issue before “counting” their preferences as part of public sector decision making. Clearly, the vast majority of the populace will not understand the complex web of factors determining the nature of species invasions, the impacts on ecosystems or production, or the nature of control options available. In such circumstances, how much weight should be given to the values “poorly informed” people place on control options? It is certainly useful to know something about how people’s WTP depends on what they know about the problem: Sheremet et al (2017), for instance, show how people’s understanding of invasive forest pathogens is related to their WTP for different management measures. Bremner and Park (2007) find a strong association between knowledge and support for invasives control programmes. This problem reflects a much more general issue in environmental economics and cost-benefit analysis when we apply the principles of economic valuation to issues such as biodiversity decline about many people will not know as much as experts (La Riviere, Czajkowski, & Hanley, 2014). Welfare economics (that part of economics underlying cost-benefit analysis and valuation) states that the economic votes (the WTP values) of everyone within the “relevant population” should count, no matter how much people know about the good in question. Thus, finding methods to help respondents understand the implications of invasive species management options before measuring their preferences, using techniques such as valuation workshops, might be viewed as a sensible approach in this regard, and economists have been investigating the ways in which deliberative mechanisms can be combined with economic valuation approaches in such situations (Lienhoop, Bartkowski, & Hansjurgens, 2015). However, these kinds of participatory approaches create aggregation problems, since now the values of observed subjects will likely differ substantially from the population from which they are drawn. Clearly, telling people more about the likely effects of an invasion, what contributes to its spread, or how the spread can be managed, will be helpful in terms of better public policy decision making and more effective management (e.g. for fire ants in Queensland: see https://www.daf.qld.gov.au/business-priorities/plants/weeds-pest-animals-ants/invasive-ants/fire-ants). But ultimately, what drives economic valuation is changes in “end-points” that make a difference to people’s well-being, whatever the nature of the complex mechanism that delivers these changes in end points. So what people care about is changes in the nature of their recreational experience in a forest, not how an invasive species produces these changes.

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The second issue to ponder is concerned with scientific uncertainty over the rate of spread of an invasive species, its impacts on ecosystems, and the effectiveness of control measures. Scientists will often be very unsure about these parameters, especially in the early stages on an invasion. From a valuation perspective, taxpayers may well care about the nature and extent of such uncertainty, in terms of their willingness to pay to support a control strategy. Sheremet el al (2017) included uncertainty over speed of spread, extent of damage and efficiency of control measures in their choice experiment on invasive forest pathogens, but found no significant effects of these levels on public WTP for a control programme (although there was a statistically significant variation with regard to how much importance people attached to uncertainty over speed of spread). More generally, however, we know that taxpayers do care about the uncertainty attached to predicted environmental policy outcomes (Lundhede, Jacobson, Hanley, Strand, & Thorsen, 2015). Thus, it seems preferable to effectively communicate scientific uncertainty over invasive species management to households and firms when trying to estimate the benefits of control. Being able to quantify this uncertainty in a way in which ordinary people can understand is a key challenge for ecologists. Third, there are issues around the treatment of irreversibility and the timing of actions. In many cases although the impacts of invasive species may be uncertain, they may also be irreversible (e.g. the extinction of an endemic species) (Finnoff, McIntosh, Shogren, Sims, & Warziniack, 2010). In these cases it has been suggested that the question should not be ‘how much are society willing to pay’ but ‘how much can society afford to lose’. Acting to prevent species invasions, or quick action of control, can prevent irreversible impacts and ensure that the widest range of options are available going forward. However the costs of prevention of irreversible changes needs to be considered. Moreover, if we can learn more about damage costs and about the effectiveness of control measures as time passes, then waiting itself generates an option value which should be considered in deciding when to act (Sims & Finnoff, 2013). While we have illustrated in this paper the economic damage caused by invasive species, in many cases invasive species also have cultural or social values, and failing to account for such values can undermine control efforts (Estévez, Anderson, Pizarro, & Burgman, 2015). In Hawaii, 12% of the population were in favour of maintaining a feral cat population, rising to 50% of people involved in animal welfare organisations. The reasons for such opinion was related to enjoyment of seeing feral cats, and an intrinsic value of knowing feral cats persist, even if not seen (Lohr & Lepczyk, 2014). Similarly in Bonaire (an island in the Caribbean), positive public attitudes towards feral donkeys restricted the possible control measures available, as any lethal control programme would be met with high social resistance (Roberts et al., 2018). Understanding the positive values associated with invasive species can be central to designing effective invasive species control measures, particularly where control is sanctioned by governing bodies who must respond to multiple competing agendas, and where invasive species control can become a politicised issue. Finally in a rapidly changing climate the very concept of invasive species becomes problematic. Huang et al (2011) illustrate that increasing temperatures are associated with increases in invasive species in the UK, USA and China, whilst Medlock and Leach (2015) show

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the increased potential of mosquito invasion into the UK under climate change. Climate change is associated with changing ranges for many species, particularly those with narrow temperature limits. While controlling increased invasive species arising from climate change may be of high importance to safeguard native species, so too is enabling species to adapt their ranges to survive a changed climate. For species where natural range expansion is limited, such as by island size or due to other environmental barriers, including urban areas or mountain ranges, translocation to novel environments may be required to ensure persistence (Braidwood, Taggart, Smith, & Andersen, 2018). Appropriate measures to prevent invasive species, and control measures to tackle invasive species, must therefore allow for, and in some cases manage, natural range expansion. 8. Conclusions We have illustrated the range of benefits of controlling invasive species, including reducing losses from agriculture, improving ecosystem health, and safeguarding biodiversity. As invasive species and their impacts continue to increase, so does the need to manage these pressures on ecosystems. Recognising the economic benefits of control, as well as the costs, provides vital information to policy makers and practitioners to prioritise invasive species control actions. Benefits of managing invasive species are not limited to those associated with market-valued goods such as crops, but should include increased exposure to disease and disruption to ecosystem service supply and impacts on wild species. Decision making also needs to account for the impact of varying knowledge on the outcomes associated with control of invasive species, for uncertainty in predicting outcomes and the potential irreversibility of not controlling invasive species. What economic valuation demands of ecologists in this regard is simple to set out, but much less simple to deliver. These demands include being able to quantify the impacts of invasives on end-points which people care about directly, or which indirectly affect their well-being over time; and the extent to which specific management actions mediate undesirable direct and indirect impacts. Given the fast-changing landscape of invasive species management, these demands are certainly not trivial in terms of the kinds of new research which needs to be undertaken. Finally, this paper has not had space to discuss the other ways (beyond the valuation of damages) in which economics can contribute to management decision-making and policy-making over invasive species. We encourage ecologists to take a look at the citations within our paper to this omitted work, where even a casual glance will show that best responses are typically determined by an interaction of ecological, epidemiological and economic parameters (MacPherson, Kleczkowski, Healey, & Hanley, 2018) . Acknowledgements: We would like to thank Paul Armsworth, Kirsty Park and the economics group at the James Hutton Institute for comments on the draft version of this paper. Braidwood, D. W., Taggart, M. A., Smith, M., & Andersen, R. (2018). Translocations,

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